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A TOXICITY INDEX FOR ASSESSING 
THE SUITABILITY OF STREAMS FOR AQUATIC LIFE 






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14 July 1988 



by 



Richard E. Sparks 

Aquatic Biology Section 

Illinois Natural History Survey 

River Research Laboratory 

Box 599 

Havana, Illinois 62644 



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Toxicity index July 14, 1988 Page 1 



A TOXICITY INDEX FOR ASSESSING THE SUITABILITY OF STREAMS FOR 

AQUATIC LIFE 

R.E. Sparks 



INTRODUCTION 

Habitat quality, water quality, and biotic interactions all affect 
aquatic organisms. Streams have been classified according to the suitability 
of the water or habitat for aquatic life (water depth, flow velocity, 
substrate, temperature, concentration of oxygen and pollutants) or according 
to characteristics of the indigenous populations (species diversity, presence 
or absence of indicator species or guilds). Although some aquatic ecologists 
consider water quality a part of habitat quality, it is useful in this paper 
to make a distinction between chemical characteristics of the water which are 
measured routinely in water quality monitoring programs and the physical 
characteristics of the habitat which are not. While population data can 
indicate that a problem exists, the cause of the problem is not always easily 
determined from available information on water quality or habitat quality. If 
the causes are unknown, it is difficult to design measures to restore the 
biological quality of the stream. The biological significance of water 
quality data is often obscure, particularly because factors are usually 
considered one at a time (does the concentration of factor X exceed the water 
quality standard for aquatic life?), whereas organisms are exposed to many 
factors simultaneously. Water chemistry is usually measured in grab samples 
taken once a month, or even less frequently, so it usually is impossible to 
determine how long organisms were exposed to stressful factors. The degree of 
stress depends not only on concentration but also on duration of exposure. 

The toxicity index described in this paper sums component toxicities 

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Toxicity index July 14, 1988 Page 2 

contributed by 20 common pollutants. The algorithms account for known effects 
of environmental factors which modify toxicity (temperature, dissolved oxygen, 
hardness or alkalinity). The toxicity index can be coupled to hydrologic and 
water quality models to estimate exposure durations as well as toxicity 
magnitudes, and to develop empirical relationships between index values, 
exposure durations, and fish community structure. 

Although toxicity indices have been developed in both North America and 
Europe, research in Illinois has contributed substantially to verification, 
refinement and novel application of toxicity indices, including application to 
stream classification. 

This paper describes: (1) the background, assumptions, and limitations of 
the toxicity index, (2) an example of the computation of the toxicity 
contributed by one component (ammonia), (3) application of the toxicity index 
to two rivers in Illinois, for the purposes of classifying reaches according 
to suitability for fish, determining which factors contribute the most 
toxicity, and for empirical determination of the relationship between the 
toxicity index and characteristics of the fish populations, and (4) future 
applications and improvements to the index. 

ASSUMPTIONS AND LIMITATIONS 

The Additive Assumption 

Sprague and Ramsay (1965), Lloyd (1965), and Sprague (1970) were the 
first to propose the toxic units approach to predicting the joint toxicity of 
mixtures of common industrial and municipal pollutants. They expressed the 
separate toxicant concentrations as fractions of their lethal threshold 
concentrations and assumed the joint lethal effects were additive. The units 
were referenced to the species used to determine the lethal thresholds in 

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Digitized by the Internet Archive 

in 2010 with funding from 

CARLI: Consortium of Academic and Research Libraries in Illinois 



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Toxicity index July 14, 1988 Page 3 

laboratory bioassays. The bluegill is the reference organism in the Illinois 
toxicity index because it is widespread in Illinois (it is the state fish) and 
is commonly used in bioassays (Lubinski 1981; Lubinski and Sparks 1981; 
Lubinski et al. 1974; Muchmore et al. 1979; and Brigham and Hey 1981). If the 
concentration of zinc in water which kills 50% of the exposed population of 
bluegill sunfish (Lepomis macrochirus) in 96 hours is 8 mg/1 (the 96-hour 
LC50), then 8 mg/1 is considered to be 1.0 toxic unit, or more specifically, 
1.0 bluegill toxicity unit .BGTU (Lubinski et al. 1974). A zinc concentration 
of 4 mg/1 in a mixture therefore is 0.5 BGTU, and is the component toxicity 
attributable to zinc. If a stream contained 0.5 BGTU of zinc and 0.5 BGTU of 
copper, the water in this stream has a toxicity index value of 1.0 (0.5 + 0.5 
= 1.0) and is predicted to be lethal to fish in 96 hours of exposure. 

Two controversies developed over this approach. One was whether 
combinations of common pollutants were in fact additive, more than additive, 
or less than additive in contributing to lethality. The other was whether the 
toxicity index could be empirically related to sublethal effects on fish 
populations, e.g., failure of reproduction or changes in species occurrence or 
dominance. 

The net results of many laboratory and field tests of the additive 
assumption are that very toxic mixtures (1.0 toxic unit) are more-than- 
additive, as measured by survival time; mixtures where component toxicities 
exceed 0.2 generally are additive, as measured by lethal thresholds; mixtures 
where component toxicities are less than 0.2 usually fail to add because the 
components apparently contribute no acute toxicity to the mixture; and 
reproducing populations of native fishes can be maintained where toxicity 
indices are less than 0.2, and habitat (cover, substrate, water depth and 
velocity) and biotic factors (food supply, predators, competitors, parasites 

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and disease organisms) are not limiting. These conclusions are supported by 
literature which will be reviewed briefly here, because of the central 
importance of the additive assumption to the toxicity index, and by the 
results of the two applications which are described later. 

The archetypical example of a more-than-additive toxic effect was 
furnished by Doudoroff in 1952. He demonstrated that Pimephales sp. could 
withstand 8.0 ppm of zinc alone or 0.2 ppm of copper alone for 8 hours, but 
the combination of only 1.0 ppm zinc with 0.025 ppm copper killed most of the 
fish in 8 hours. This example has been cited frequently and used as a warning 
of the type of more-than-additive effects to be expected if more and more 
pollutants are introduced to the environment, without an assessment of their 
joint effects (Cairns 1957). More-than-additive effects of zinc-copper 
mixtures have also been demonstrated by other investigators (Lloyd 1961b). 

It is therefore surprising that Herbert stated in 1965 that no cases had 
been found in which the toxicity of mixtures of poisons commonly found in 
sewage and industrial wastes (copper, zinc, lead, phenol, ammonia, and 
cyanide) were appreciably more than additive. This apparent contradiction is 
resolved by looking more closely at the concentrations and exposure times used 
in the experiments. As Sprague and Ramsay (1965) pointed out, survival times 
in mixtures where the total toxicity appreciably exceeds 1.0 toxic unit 
(Doudoroff's example) are shorter than expected. When lethal thresholds (96- 
hr LC50's or LT50's) are measured, the toxicities of mixtures of heavy metals 
and other toxicants generally do add up in laboratory experiments. However, 
there are exceptions as shown below. 

Sprague (1970) pointed out that the toxicity of mixtures of phenol, 
ammonia, and zinc was overestimated by Brown et al. (1969) in three out of 
four cases where two of the toxicants were present at 0.10 to 0.14 toxic 

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units. When each toxicant was present at 0.2 toxic units or more, the 
toxicities did add up. These data suggest that the threshold for acute toxic 
effect postulated by Lloyd (1965) does exist, at least for metals, and that it 
may be approximately 0.2 toxic unit. 

Results using mixtures other than metals alone are ambiguous. In 
Illinois Sparks and Anderson (1977) reported that a toxicity index 
underestimated the lethality of a mixture of linear alkyl sulfonate (LAS, a 
detergent), ammonia, and zinc by approximately 50%. In contrast, Esvelt et 
al. (1971) accurately estimated the toxicity of wastes entering San Francisco 
Bay by adding the toxic contributions of methylene blue-active substances 
(MBAS, mostly detergents) and ammonia. Herbert (1962) noted an 82.5 percent 
agreement between predicted and observed toxicities in field tests, although 
Sprague (1970) noted that the limit set for agreement was rather wide. In 
fresh-water reaches of four rivers in England, actual 48-hr LC50s approximated 
65% of predicted values, and in the saline estuaries, prediction further 
underestimated actual toxicity (Brown et al. 1970). Lloyd and Jordan (1964) 
found their index consistently underestimated the toxicity of sewage effluents 
and that the relation between the predicted and observed toxicity was 
described by the function: 

y = 1.25x - 0.59 
where y is the observed toxicity and x the predicted toxicity. 

Some of the underestimates of toxicity reported in the older literature 
probably result from bioassays which were run for arbitrary time periods, 
e.g., 8-, 24-, or 48-hours, which were too brief for full uptake of the 
toxicant and full exertion of the toxic effect in the test populations, 
particularly at the lower concentrations. The individuals in every test 
population differ in tolerance to the toxicant, and the difference is 

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expressed in survival time. The thresholds estimated from short-term 
bioassays could underestimate actual thresholds. Modern practice is to 
continue an acute bioassay for at least 96 hours, or better, until all 
mortality (or other toxic effect) has ceased for at least 24 hours. 

Even with these limitations of the additive assumption however, most of 
the predicted toxicities of mixtures differed from the actual toxicities by no 
more than 50%. Considering that the modes of action of many of the toxicants 
are unknown, let alone the modes of interaction with other toxicants and with 
modifying factors in the water or in the organisms, it is remarkable that the 
error is not higher. It is likely that the error will be reduced as more is 
learned about interactions, which then can be incorporated in the toxicity 
index. Acceptance of a 50% error appears preferable to the alternative of not 
using the existing bioassay literature and chemical monitoring data to 
estimate the joint effects of toxicants in lakes and streams. In cases such 
as spills of complex wastes, where the predicted toxicity is several times 
greater than 1.0, a 50% error is insignificant. At the other extreme, field 
evidence from a variety of polluted rivers in England and Illinois supports a 
relatively narrow range of 0.2-0.4 for a threshold below which populations of 
native fishes can sustain themselves, if other factors are not limiting (this 
paper; Brigham and Hey 1981; Herbert et al. 1965; and Brown 1970). 

Bioaccumulation 

By definition, bluegill toxicity indices sum 96-hour LC50 values and thus 
pertain only to acute toxicity, or to empirically determined relationships 
between index values and fish populations in streams. However, similar 
indices can be based on chronic effects when sufficient data are developed. 
The index does not assess effects of toxicants which accumulate in organisms 

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from levels in water which are too low to have acute effects. The index 
therefore does not identify chemicals which might not affect fish but might 
affect consumers of fish. 

Choice of Species and Life Stage 

The present toxicity index is based on juvenile and adult life stages of 
the bluegill because most of the available toxicological information was for 
these stages. Additional toxicity indices, however, could be based on 
sensitive embryonic and larval stages of organisms (e.g. Reinbold and 
Pescitelli 1982) representing a variety of trophic levels or functional groups 
(e.g. Anderson et al. 1978) within aquatic systems. Multiple indices would be 
better predictors of species replacements or ecosystem-level effects. 

Limitations of the Data Bases 

Algorithms for computing component toxicities can be continually updated 
as more toxicity data become available on newly introduced chemicals and on 
interactions and modifying factors, but in the meantime, there are data gaps. 
For example, I could locate no information on the toxicity of fluoride to 
bluegills. Neuhold and Sigler (1960), however, reported 96-hr LC50s for 
fluoride-sensitive rainbow trout and fluoride-tolerant common carp, and it was 
reasonable to assume bluegill would be intermediate in sensitivity. 

Automated samplers, event-triggered sampling (during a flood, drought, or 
spill), and water quality models (calibrated with available data) can overcome 
the limitations imposed by the relatively infrequent sampling (usually once a 
month, at best) characteristic of most water quality monitoring programs. It 
is particularly important to know the duration of exposure because fish can 
survive brief exposures to conditions (low oxygen, lethal concentration of a 
toxicant) which eventually would kill them. Another limitation of the water 

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quality data base is that total concentrations are measured, rather than toxic 
fractions. Doudoroff et al. (1966) demonstrated that molecular cyanide (HCN) 
is the toxic agent in solutions containing ionized cyanide (CN~) and cyanide- 
metal complexes, but only total cyanide is measured by IEPA and it may not be 
reliable to compute the molecular fraction of complex cyanide solutions in 
streams, even when the general chemical composition and pH are known 
(Doudoroff 1976). If nontoxic cyanide complexes are present, use of the total 
cyanide concentration overestimates the toxicity. These limitations could be 
overcome with better analytical techniques or more sophisticated programs for 
calculating chemical equilibria in complex mixtures. 

CALCULATION OF COMPONENT TOXICITIES: AN EXAMPLE USING AMMONIA 

For some toxicants, the effects of environmental factors on the chemical 
form and concentration of the toxicant, and on the resistance of the aquatic 
organism, have been determined in the laboratory and can be accounted for in 
the toxicity index. An example, using ammonia, is described next. 

Ammonia is a common pollutant in Illinois waters. It is formed by the 
breakdown of urea, and hence is found in effluents from livestock confinement 
areas and sewage treatment plants. Ammonia, in several chemical forms, is 
stored, transported by pipe, truck, rail and barge, and applied to 
agricultural lands as a nitrogen source for crops. It occurs in effluents 
from refineries and munitions industries. 

Effects of Modifying Factors on Chemical Equilibria 

Ammonia in water exists in a toxic, un-ionized form, NHo, and a non-toxic 
ionized form, NH^ + , with the equilibrium between the two forms governed by pH, 
temperature, and salinity, although in freshwater the salinity effect can be 

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ignored (Emerson et al. 1975). The total ammonia-nitrogen concentration 

(ionized + un-ionized) is measured in most water quality monitoring networks, 

but the proportion which exists in the toxic state can be determined from the 

field pH and temperature, using the tables or two equations provided by 

Emerson et al. (1975: 2382). 

NHj-N = total ammonia-N x 1 

1 + antilog (pka -pH) 

where pka = the negative log of the ionization constant: 

pka = -0.03229 (temp °C) + 10.05333 



Effects of Modifying Factors on Sensitivity of the Fish 

The toxicity of un-ionized ammonia to fish varies with the size of the 
fish and the temperature and dissolved oxygen concentration of the water 
(Roseboom and Richey 1977; Reinbold and Pescitelli 1981; Merkens and Downing 
1957). The toxicity of ammonia increases at low temperatures probably because 
the overall metabolic rate of the fish is lower, and their ability to excrete 
ammonia in their urine is reduced. Fish pick up ammonia from the water via 
their gills, and form ammonia within their tissues as a waste product of 
protein metabolism. When the rate of ammonia uptake and production exceeds 
the rate of ammonia excretion, the internal concentration of ammonia rises to 
lethal levels (Fromm 1970; Brockway 1950). Small fish are more sensitive to 
ammonia than large fish, presumably because the ratio of gill surface 
available for ammonia uptake to body volume or mass is greater in the smaller 
fish. The toxicity of every chemical used in the toxicity index, including 
ammonia, increases at low dissolved oxygen concentrations, presumably because 

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low oxygen itself is a stressor, and virtually any reduction below saturation 
reduces the metabolic scope for excretion or detoxification. As ambient 
oxygen concentrations decline, some fishes can compensate to some extent by 
increased ventilation frequency or volume (Marvin and Heath 1968), but 
probably at the expense of other metabolic activities, including ammonia 
excretion. 

The 96-hour LC50 for un-ionized ammonia (as NH^-N) was regressed against 
fish weight (fwt) and water temperature (temp), using Illinois data of 
Reinbold and Pescitelli (1981) and Roseboom and Richey (1977). The lowest and 
highest temperatures used by these investigators were 4 and 28°C, and the 
regression equation is not extrapolated beyond the range of these data: 

if temp < 4 °C, then the 96-hour LC50 = i (° 26639 fwt + -025353(4) - .67645) 
if temp > 28 °C, then the 96-hour LC50 = i (- 026639 fwt + -025353(28) - .67645) 
otherwise, the 96-hour LC50 = i (- 026639 fwt + •025353(temp) - .67645) 
Next, the LC50 is adjusted to reflect the increased toxicity of un- 
ionized ammonia at dissolved oxygen (DO) concentrations below saturation 
(Merkens and Downing 1957): 

LC50 = LC50 (at 100% saturation) x .013297 (DO, % saturation) - .32965 
Unfortunately, only 2 levels of dissolved oxygen were tested. I assumed a 
linear relationship between the LC50 and DO. 

Application to an Ammonia Spill in the Illinois River 

For the purposes of this example, fish weight is set at 1 gram, which 
would probably be at the lower end of the average weight for bluegills in 
their first year of life in January (Carlander 1977) when 622,000 gallons of 
urea ammonium nitrate solution spilled into the Illinois River at Seneca, 
Illinois. A dispersion model indicated that the total ammonia concentration 

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Toxicity index July 14, 1988 Page 11 

6.5 river miles downstream, at Marseilles, was 47 mg/1 for approximately 19 
hours (personal communication, 8 April 1988, Mr. Thomas Butts, Professional 
Scientist and Assistant Head, Water Quality Section, Illinois State Water 
Survey, Peoria, Illinois). 

Sensitivity of Ammonia Component Toxicity to Modifying Factors. 
Dissolved oxygen levels have a marked effect on the predicted toxicity of 47 
mg/1 total ammonia-N (Table 1). At saturation, the ammonia spill exceeded 
twice the lethal concentration, but at 13% saturation, the component toxicity 
was greater by 3 orders of magnitude, and would probably have killed fish 
within a few hours at summer temperatures. At colder temperatures, less of 
the total ammonia exists in the toxic, un-ionized form, but the sensitivity of 
the fish increases, so the net change in component toxicity is rather small, 
from 2.69 to 1.90, as the temperature declines from 28 ° C to 4 ° C (Table 1). 
At cold temperatures, fish may be exposed to ammonia longer than it takes a 
spill to pass a fixed point, because they lose their equilibrium and float 
upside down (Reinbold and Pescitelli 1981) and thus would be carried along in 
the spill. 

Conclusions 

Even assuming the DO was at saturation, and no other toxicants were 
present, the toxicity attributable to the spill was approximately twice the 
lethal threshold at the cold temperatures expected in January. Fish which 
lost their equilibrium were exposed to these lethal concentrations longer than 
19 hours, and very probably died. 

The component toxicity approach was a rational way of integrating 
information on environmental factors, chemical concentrations, and 
susceptibility of organisms to assess the probable degree of harm to aquatic 
life caused by this ammonia spill. The next application examines toxicities 

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contributed by 20 components in the Illinois River and the connecting Chicago- 
area canals, collectively known as the Illinois Waterway. 

APPLICATION TO THE ILLINOIS WATERWAY 

The Illinois Waterway extends from the highly altered rivers and 
waterways in Chicago downstream 326 miles via the Illinois River to the 
Mississippi River upstream of St. Louis. The numbers and kinds of native 
gamefish and their condition declines in the upstream direction, towards 
Chicago, where fish populations are dominated by a few species, including the 
introduced carp and goldfish (Sparks and Starrett 1975). 

Procedures 

In the years 1972-1974, the toxicity index was applied to the Illinois 
Waterway, to determine whether high toxicity values were associated with the 
degraded fish populations in the upstream reaches, and to determine what 
components contributed the most toxicity. Field tests, using bluegills 
exposed to river water for 4 days in cages in the river and in aerated plastic 
pools on shore (where the water was renewed by pumping), were conducted at an 
upstream (Dresden) and a downstream (Beardstown) site in 1974 (Lubinski et al. 
1974). Controls were maintained nearby in cages in less polluted tributaries. 

Water quality data were obtained from the Illinois Environmental 
Protection Agency (IEPA) for samples taken 4 to 13 times per year at 20 
stations along the Illinois Waterway. The same data were obtained by the 
Natural History Survey on samples taken at least once daily during the field 
tests. 



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Results 

Field bioassays. The field bioassays confirmed that no short-term 
mortalities attributable to toxicity occurred in the two reaches of the 
Illinois River or in the tributaries where the daily index values were less 
than 0.14 (Lubinski et al. 1974). No mortalities occurred at Dresden. At the 
Beardstown site, at the mouth of the Sangamon River, there were no mortalities 
in the aerated swimming pool receiving Illinois River water, but 22% of the 50 
test fish confined in the cage in the Illinois River died, probably because of 
the stress of swimming against a relatively high current velocity (0.4-0.6 
m/sec) or being forced against the mesh by the current (Lubinski et al. 1974). 
Twelve percent of the fish died in a backwater area receiving flow from the 
Sangamon River where the water levels were falling and the cage had to be 
lifted from the mud and moved to deeper water several times, stressing the 
fish. 

Mean Toxicity Indices. The mean annual toxicity indices in the Illinois 
Waterway (the sum of the toxicity indices at each station, divided by the 
number of samples taken that year) generally were below 0.1 at the station 
farthest upstream, which receives clean water from Lake Michigan via a 
navigation lock, well above 0.2 through the Chicago waterways, and 0.1 or less 
starting where the Chicago Sanitary and Ship Canal joins the Des Plaines River 
(Figure 1). The year 1972 was an exception, with elevated toxicity indices at 
Peoria. 

Maximum Toxicity Indices. Extreme events inevitably go undetected 
because of the small number of water quality samples taken per year by the U. 
S. Geological Survey and the Illinois Environmental Protection Agency (12 or 
less). If it takes approximately 5 minutes to fill the sample bottle each 
month, there are 43,195 minutes in which no samples were taken (60 minutes x 

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24 hours x 30 days = 43,200, less 5 minutes = 43,195). The sample represents 
conditions occurring during only 1 / 10,000th of the month. Maximum annual 
toxicity indices partially compensate for this limitation by adding the peak 
toxicities contributed by each component during the year. The assumption is 
that the highest component toxicities occurred simultaneously, and the maximum 
index approximates "worst case" conditions. Such an assumption is not 
completely unreasonable because: (1) if one toxicant is at high concentration 
because of an excessive discharge, several others usually are too, because 
industrial wastes typically are complex mixtures of pollutants, and (2) during 
low flow conditions, all wastes are more concentrated because there is less 
dilution capacity. (A better, but more costly simulation procedure, in terms 
of data and programming requirements, which does not make the "simultaneous 
maxima" assumption, is described in application 2 below.) Maximum toxicity 
indices approach 0.5 in the upper Illinois Waterway (except for the uppermost 
station, near Lake Michigan), and generally decline downriver, with the 
exception of the Peoria stations in 1972 (Figure 1). 

Component Toxicity: Cyanide. The 1972 peak in toxicity at Peoria is 
explained largely by the cyanide component (Figure 2), with some contribution 
from zinc and copper (relatively high values, compared to other years, did 
occur together in one sample in this case). The spiky, highly variable 
cyanide pattern is typical of toxicants which are accidentally or sporadically 
introduced, and it is fortuitous that the monthly sampling in 1972 happened to 
detect an apparent spill which was moving downstream, and which probably 
originated from an industrial source in Peoria. 

Component Toxicity: Ammonia. The dominant contributor to the pattern of 
high toxicity in the Chicago waterways and declining toxicity downstream is 
ammonia (Figure 2), which is continuously released from the sewage treatment 

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plants in Chicago. Ammonia probably declines downstream because of dilution 
by tributaries, uptake by aquatic plants in the upper Illinois Waterway, and 
conversion to nitrate by bacteria in the water. Toxicity in portions of the 
Illinois Waterway in Chicago could be reduced below an index of 0.2 by 
increased diversion of clean dilution water from Lake Michigan, but the 
diversion increase would have some potentially deleterious effects, including 
possible scouring of toxicant-laden sediments from the waterways downstream 
into more biologically productive areas (Havera et al. 1980). 

Despite improvements in sewage treatment in the Chicago River since the 
early 1970s (Macaitis et al. 1987), ammonia remains a problem in much of the 
Illinois River. Unpublished data from the U. S. Fish and Wildlife Service 
(personal communication, 29 February 1988, Richard Ruelle, Aquatic 
Toxicologist, Environmental Services Section, U. S. Fish and Wildlife Service 
Field Office, Rock Island, Illinois) indicates that sediments in backwaters at 
least as far as 200 miles downstream from Chicago contain ammonia which is 
released in sufficient quantities when the sediments are agitated to kill fish 
in 96-hour laboratory bioassays. Sediments are frequently agitated by wind- 
and boat-generated waves (Jackson and Starrett 1959; Sparks and Starrett 
1975). Sparks (1984) reported that un-ionized ammonia concentrations in 
excess of the lethal level for fingernail clams (0.06 mg/1 NH^-N) occurred in 
the Illinois River at Marseilles, Hennepin and Lacon in 1980, probably because 
uptake of CC»2 and HCO3- during algal blooms raised the pH above 8 and 
increased the proportion of ammonia existing in the toxic un-ionized form. 
Fingernail clams are an important food for diving ducks and bottom-feeding 
fish, both of which were adversely affected by the die-off of the clams and 
other mud-burrowing invertebrates in a 100-mile reach of the Illinois River in 
1958 and by their continuing failure to recolonize (Mills et al. 1966; Sparks 

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Toxicity index July 14, 1988 Page 16 

1984). 

Summary of Application 1 

In summary, the toxicity index proved useful in interpreting the water 
quality data collected by the IEPA on the Illinois Waterway to determine what 
toxicants are acutely limiting to fish and should be controlled to restore 
fish populations. Index values above 0.2 in the upper Illinois Waterway are 
associated with depauperate fish populations in generally poor body condition. 
The toxicity index falls below 0.2 downstream, where more species of native 
fishes, in better condition, occur. The principal contributor to the total 
acute toxicity is ammonia, with sporadic contributions from cyanide. Ammonia 
originates from such a widely dispersed, large capacity sewage system 
(Chicago) that pulses are absorbed and ammonia loading is relatively constant, 
but downstream variations in temperature, pH, and dissolved oxygen alter the 
component and overall toxicities. Therefore, an important limitation was the 
infrequency of water quality sampling (only 4 to 13 times per year), which 
made it difficult to detect extreme conditions or determine how long fish were 
exposed. The next application describes how water quality modeling can be 
used to overcome this limitation and, when coupled with the toxicity index, 
used to predict the effects of alternative pollution control measures on fish 
populations. 

APPLICATION 2: EVALUATING STRATEGIES FOR WATER QUALITY MANAGEMENT 

Objectives 

The objectives of this project were (1) to explore the relationship 
between fish communities and their physical and chemical environments, using 
the good water quality and fish population data sets and a calibrated water 
quality model available for the DuPage River in northeastern Illinois, (2) to 

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develop a continuous toxicity function for relating the status of these 
communities to spatial and temporal patterns of physical and chemical events, 
and (3) to use the resulting toxicity functions to predict effects of 
alternative pollution control measures on fish populations (Brigham and Hey 
1981). 

Procedure 

Three stream reaches of the DuPage River were selected for modeling and 
analysis, based on known differences in fish faunas. One reach supported a 
mixed community of 17 species, including bluegill and carp, (hereafter 
referred to as the bluegill reach), one supported only 6 species, excluding 
bluegill and dominated by carp (carp reach), and the third reach was fishless 
(Brigham and Hey 1981). The toxicity index was calculated for each reach at 
1-hour intervals over a simulated span of 3 years, using water quality values 
generated by hydrologic and water quality models which were calibrated for the 
reaches. The models were developed by Hydrocomp, Inc., Palo Alto, California 
and implemented by the Northeastern Illinois Planning Commission. The 
toxicants were un-ionized ammonia, cyanide, lead, zinc, copper, linear 
alkyl sulfonate detergents (LAS), and total residual chlorine (TRC). After 
initially observing the magnitude and variability of the toxicity function 

generated for each reach (Figure 3), certain water quality input variables 

were modified to simulate different management practices on each stream reach. 

Results 

Critical Thresholds. The toxicity indices for the three stream reaches 
on 17 March 1972 are typical of the run simulating the period from 1 October 
1970 to 30 September 1973 (Figure 3). Toxicity levels regularly exhibited in 

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these reaches compared well with the results of previous investigators, who 
indicated that the effects of acute toxicity in altering the species 
composition of fish communities became measurable in streams at levels of 0.2 
to 0.4 toxic units (Lloyd and Jordan 1964; Edwards and Brown 1966). 

Relationships between Frequency, Duration and Intensity of Exposure and 
Fish Populations. Perhaps more important than the typical values however, are 
the frequency and duration of episodes where toxicities exceeded the lethal 
value of 1.0 and the empirically determined threshold of 0.2-0.4 (Table 2). In 
the fishless reach , the toxicity index made 194 excursions above 3.0 lasting 1 
hour or more during the 3 years, 39 excursions lasting 24 hours or more, and 
21 lasting 96 hours or more. The carp reach exhibited toxicity levels of 1.0 
unit for 96 hours or more on 14 occasions, whereas the bluegill reach never 
exceeded 0.3 for even 1 hour. There were only 2 occasions during the 
simulated 3-year period when the toxicity index in the bluegill reach was 
between 0.25 and 0.30 for 24 hours or more. The average length of these 2 
excursions was 36.5 hours (Brigham and Hey 1981). 

Simulation of Management Alternatives 

Assume a management goal of changing the fishless reach to a reach 
capable of supporting fish. A typical management plan might include: (1) 
reduction of ammonia concentrations from wastewater treatment plants to 1.5 
mg/1 during the summer and 4.0 mg/1 during winter, (2) elimination of combined 
(stormwater and sewage) sewer overflows, (3) reduction of sediment oxygen 
demand, and (4) moderate increase of dissolved oxygen in the wastewater 
effluents or in the stream (Brigham and Hey 1981). 

This plan primarily targets ammonia toxicity, which is reduced by 3 
orders of magnitude, from 20.2 to 0.023 (Table 3). The mean toxicity index 
declines by an order of magnitude, from 23.0 to 2.12 (Table 3), but still 

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significantly exceeds the mean index of 1.04 in the carp reach (Brigham and 
Hey 1981). The largest remaining contributor to the total toxicity is 
chlorine (mean component toxicity = 1.94). Cessation of effluent chlorination 
would reduce the mean index to 0.186, which is very close to the mean of 0.115 
in the bluegill reach. If no excursions above 0.3 occurred, and no factors 
other than toxicity are limiting, a mixed community of native fishes probably 
could be maintained. 

This simulation was run in 1981, and the Illinois Environmental 
Protection Agency has since abandoned the requirement for effluent 
chlorination, based on evidence that there would be little additional public 
health risk from infectious diseases and much improvement in water quality for 
aquatic life. The latter evidence included studies which employed or referred 
to the toxicity index (Muchmore et al. 1979; Dreher 1981; Hey, Pappas and Cox 
1980; and Hey et al. 1982). Fish populations in the Chicago waterways have 
shown recent improvement following discontinuation of effluent chlorination 
(personal communication, 1 March 1988, Mr. Samuel Dennison, Fisheries 
Biologist, Metropolitan Sanitary District of Greater Chicago). 

FUTURE DEVELOPMENT 

The toxicological data base for the index was last revised in March 1981. 
The programs for computing toxicity indices from IEPA water quality data are 
written in BASIC for a Tektronix 4051 microcomputer and a CYBER 175 at the 
University of Illinois (Lubinski 1981). The University is replacing the CYBER 
and the Tektronix 4051s are no longer in common use. The algorithms should be 
rewritten, using updated data, for IBM- or Apple-compatible personal 
computers, and a new user's guide prepared. 

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The revised index then should be applied to reaches intermediate in 
toxicity between the bluegill reach and the carp reach of the DuPage River 
study (Brigham and Hey 1981) to quantify more precisely the timing, frequency, 
duration, and intensity of exposures which cause shifts in community 
structure. The differences in toxicity and exposure patterns between the 
bluegill and carp reaches on the DuPage River were too great to determine, for 
example, whether the community structure characteristic of the bluegill reach 
would degrade if the index exceeded 0.3 for brief periods. 

This type of information is useful in pollution control engineering, 
where a performance standard is achieved within some specified variation and 
failure rate. If the biological consequences of excursions beyond the mean 
can be specified, then waste control programs can be designed to stay within 
the limits without incurring unnecessary expense to achieve lower variation or 
failure rates. 

The timing of excursions also should be examined, e.g. do excursions in 
the spring when larval fish are present have greater effect on fish 
populations than the same excursions in late summer? If so, the waste loading 
could be adjusted seasonally to protect aquatic life in the stream. Although 
the present index, which is based on toxicity to adult fish, can be related 
empirically to the status of fish populations in streams, as described in the 
above applications, another approach would be to develop toxicity indices for 
sensitive life history stages and use them at the appropriate season. 

The present index is based on concentrations of toxicants in water. In 
the Illinois River, water quality has improved without a concomitant recovery 
of infaunal macroinvertebrate communities, because of an apparent legacy of 
toxicants remaining in the sediments (Sparks 1984). Toxicity indices should 
be developed for reference species representing several trophic levels and 

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functional groups, used together to increase the reliability of simulations 
and estimations, and verified by field trials or by application to data-rich 
environments, as described in this paper. 

In particular, indices should be developed for assessing the quality of 
sediments, as well as quality of water, using benthic macroinvertebrates and 
rooted aquatic macrophytes as reference organisms. Sediment LC50s (in mg of 
toxicant per kg of "standard" sediment materials, e.g. montmorillonite clay or 
natural sediments of consistent composition) could be determined by adding 
reagent grade toxicants to a sediment slurry, allowing it to settle, then 
adding the test organisms. The additive assumption should be tested with 
mixtures of toxicants in sediments. The database on sediment LC50s would be 
employed in a sediment toxicity index just as described above for the water- 
based index. 

SUMMARY 

The toxicity index provides a way of relating water quality monitoring 
data to toxicity data available in the literature and to fish populations in 
streams, so that stream reaches can be classified according to their 
suitability for fish communities of varying sensitivity to common pollutants 
and environmental stressors. The index goes beyond classification however, to 
identification of causative factors. Chemical concentration units are 
converted to toxicity units, so the chemicals which contribute the most 
toxicity in a stream reach can be identified. The component toxicities also 
can be summed to provide an estimate of the total toxicity in a reach. The 
assumption of additive effects appears generally valid if lethal thresholds, 
rather than survival times, are used to measure toxicity. Predicted lethal 
thresholds are generally within 50% of measured thresholds in laboratory and 

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field experiments where complex mixtures are present. 

Where sufficient data exist, the effects of factors which modify chemical 
equilibria or the sensitivity of fish (temperature, pH, dissolved oxygen, 
calcium concentration or hardness) can be taken into account in the 
algorithms. One indirect, beneficial result of this systematic search for 
information on interactions is that toxicological data gaps and research needs 
are identified and prioritized. 

Application of the toxicity index to 1970s water quality data from the 
Illinois River indicates that ammonia from the Chicago area is a major 
contributor to toxicity in the upper river, with sporadic contributions from 
cyanide downstream of Peoria. Mean index values above 0.2 occur in upstream 
reaches where there are depauperate fish communities dominated by introduced 
carp and goldfish. Native fish increase downstream where mean and "worst 
case" index values are generally below 0.2. The index also was used in an 
after-the-fact analysis to determine that a 1988 spill of urea ammonium 
nitrate into the upper Illinois River probably killed fish several miles 
downstream. 

Toxicity indices in the range 0.2-0.4 have been established as the 
threshold for alteration of the species composition of fish communities by 
results of field tests in the Illinois River and in several English rivers and 
by analysis of toxicity simulations in 3 reaches of the Dupage River: a 
fishless reach, a carp-dominated reach, and a reach with a mixed community 
including the native bluegill sunfish. A time dimension was added to this 
threshold in the DuPage River study, where the index never exceeded a 
sublethal value of 0.30 for even 1 hour in the bluegill reach, there were only 
2 excursions between 0.25 and 0.30 (lasting an average of 36.5 hours), and the 
average toxicity was 0.115 during a simulated 3-year period. 

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Toxicity indices are not designed to assess bioconcentration effects or 
to be used in place of direct in-plant or in-stream toxicity testing programs. 
The projects described here, however, have demonstrated that these indices can 
be used to complement water quality monitoring programs by providing numerical 
values that describe a biological parameter (toxicity). The recent regulatory 
emphasis that has been placed on effluent toxicity testing and biological 
monitoring suggests that the results of water quality monitoring programs are 
of limited value in assessing toxicity problems. Although it is true that a 
limited number of in-stream measurements for a particular toxicant should not 
be used in sole support of any important water management decision, the 
alternative of not considering these data at all seems equally unacceptable 
and, in fact, undermines a common objective of water quality monitoring 
programs, which is to provide information on which to base management 
decisions. 

Unfortunately, the products of most water quality monitoring programs are 
voluminous tables of data, which decision makers find difficult to interpret. 
Analysis usually is confined to the number of times standards for individual 
constituents were exceeded rather than to interactions and their biological 
consequences. Toxicity indices provide a logical way to assess the joint 
action of toxicants and the modifying effects of environmental factors on 
aquatic organisms. 

As demonstrated by the projects described here, toxicity indices can be 
used to determine which chemical components contribute the most toxicity at a 
given location or time, to relate temporal and spatial variations in water 
quality to fish community structure, to evaluate alternative pollution control 
strategies, to assess the biological effects of spills, and to classify stream 
reaches according to their suitability for fish. 

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ACKNOWLEDGMENTS 

The Illinois River application was supported by the Office of Water 
Research and Technology, Project No. A-067-ILL, and by the Illinois Institute 
for Natural Resources, Project No. 20.107. Kevin Anderson and Yip Tai-Sang 
wrote the computer programs for this project. 

The DuPage River application was supported by the U.S. Environmental 
Protection Agency, Grant No. R805614010, and was conducted jointly by Dr. 
Warren Brigham, Illinois Natural History Survey, and Donald Hey, consultant to 
the Northeastern Illinois Planning Commission and Hydrocomp, Inc. 

Dr. Kenneth S. Lubinski, currently with the U. S. Fish and Wildlife 
Service in LaCrosse, Wisconsin, contributed substantially to various projects 
involving toxicity indices from 1973 to 1974, and again from 1979 to 1987. 

K. Douglas Blodgett, Illinois Natural History Survey, River Research 
Laboratory, Havana, wrote a program in LOTUS to calculate ammonia component 
toxicity, using updated information, for analysis of the 3 January 1988 
ammonia spill in the Illinois River. 



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LITERATURE CITED 

Anderson, K.B., R.E. Sparks, and A. A. Paparo. 1978. Rapid assessment of 

water quality using the fingernail clam, Musculium transversum. 

Illinois Water Resources Center Report No. 133, University of 

Illinois. 1 15 p. 
Brigham, W., and D. Hey. 1981. A stress function for evaluating 

strategies for water quality management. Contract Report. U.S. 

Environmental Protection Agency Grant No. R805614010. 92 pp. 
Brockway, D. 1950. Metabolic products and their effects. Progressive 

Fish Culturist 12:127-129. 
Brown, V.M., D.H.M. Jordan, and B.A. Tiller. 1969. The acute toxicity 

to rainbow trout of fluctuating concentrations and mixtures of 

ammonia phenol and zinc. Journal of Fisheries Biology 1:1-9. 
Brown, V. M, D. G. Shurben, and D. Shaw. 1970. Studies on water 

quality and the absence of fish from some polluted English rivers. 

Water Research 4:363-382. 
Cairns, J., Jr. 1957. Environment and time in fish toxicity. 

Industrial Wastes 1:1-15. 
Carlander, K. D. 1977. Handbook of freshwater fishery biology. Vol. 2. 

The Iowa State University Press, Ames, Iowa. 431 p. 
Doudoroff, P. 1976. Toxicity to fish of cyanides and related compounds, 

a review. U. S. Environmental Protection Agency Ecological Research 

Series No. EPA-600/3-76-038. 155 p. 
Doudoroff, P., G. Leduc, and C.R. Schneider. 1966. Acute toxicity to 

fish of solutions containing complex metal cyanides, in relation to 

concentrations of molecular hydrocyanic acid. Transactions of the 

American Fisheries Society 95:6-22. 

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Dreher, D.W. 1981. Study of fish toxicity in the East Branch DuPage 
River. Report. Northeastern Illinois Planning Commission. 

Edwards, R.W., and V.M. Brown. 1966. Pollution and fisheries: a 

progress report. Water Pollution Control 66:63-78. 
Emerson, K., R.C. Russo, R.E. Lund, and R>V> Thurston. 1975. Aqueous 

ammonia equilibrium calculations: effect of pH and temperature. 

Journal of the Fisheries Research Board of Canada 32:2379-2383. 
Esvelt, L. A., W. J. Kaufman, and R. E. Selleck. 1971. Toxicity removal 

from municipal wastewaters. SERL Report No. 71-1. Sanitary 

Engineering Research Laboratory, College of Engineering and School 

of Public Health, University of California at Berkeley. 224 p. 
Fromm, P.O. 1970. Effect of ammonia on trout and goldfish. Pages 9-22 

in Toxic Action of Water Soluble Pollutants on Freshwater Fish. 

Report No. 18050 DST 12/70. U. S. Environmental Protection Agency, 

Water Quality Office, Washington, D.C. 
Havera, S.P., F.C. Bellrose, H.K. Archer, F.L. Paveglio, Jr., D.W. 

Steffeck, K.S. Lubinski, R.E. Sparks, W.U. Brigham, L. Coutant, S. 

Waite, and D. McCormick. 1980. Projected effects of increased 

diversion of Lake Michigan water on the environment of the Illinois 

River valley. U.S. Army Corps of Engineers, Chicago District. 550 

P- 

Herbert, D.W.M. 1962. The toxicity to rainbow trout of spent still 
liquors from the distillation of coal. Annals of Applied Biology 
50:755-777. 

Herbert, D.W.M, D.H.M. Jordan, and R. Lloyd. 1965. A study of some 
fishless rivers in the industrial Midlands. Journal of the 

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Toxicity index July 14, 1988 Page 27 

Proceedings of the Institute of Sewage Purification 6:569-582. 
Hey, D.L., E.L. Hardin, D.W. Dreher, and N.S. Philippi. 1982. Proposed 

revision to the water quality standards for the DuPage River. 

Report. Northeastern Illinois Planning Commission. 88 p. 
Hey, D.L., J.M. Pappas, and L.C. Cox. 1980. An economic analysis of 

effluent standards for BOD, ammonia, total suspended solids, and 

disinfection: case study of a modern treatment plant. Document No. 

80/25. Illinois Institute of Natural Resources, Environmental 

Management Division, Chicago. 46 p. 
Jackson, H.O., and W.C. Starrett. 1959. Turbidity and sedimentation at 

Lake Chautauqua, Illinois. Journal of Wildlife Management 23:157- 

168. 
Lloyd, R. 1961. The toxicity of mixture of zinc and copper sulphates to 

rainbow trout (Salmo gairdnerii Richardson). Annals of Applied 

Biology 49:535-538. 
Lloyd, R. 1965. Factors that affect the tolerance of fish to heavy 

metal poisoning. Pages 181-187 in C. M. Tarzwell, ed. Biological 

problems in water pollution, third seminar. U.S. Department of 

Health, Education, and Welfare, Public Health Service, Division of 

Water Supply and Pollution Control, Cincinnati, Ohio. 
Lloyd, R., and D.H.M. Jordan. 1964. Predicted and observed toxicities 

of several sewage effluents to rainbow trout: a further study. 

Journal of the Proceedings of the Institute of Sewage Purification, 

Pt. 2, pp. 183-186. 
Lubinski, K.S. 1981. Modification of a bluegill toxicity index system 

for use by the Illinois Environmental Protection Agency: Phase II. 

Bluegill toxicity index systems description and protocol 

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Toxicity index July 14, 1988 Page 28 

development. Contract Report to the Illinois Environmental 

Protection Agency, Springfield. 80 p. 
Lubinski, K.S. and R. E. Sparks. 1981. Use of bluegill toxicity indexes 

in Illinois. Pp. 324-337 in D.R. Branson and K.L. Dickson, ed. 

Aquatic Toxicology and Hazard Assessment: Fourth Conference. ASTM 

Special Technical Publication No. 737. American Society for Testing 

and Materials, Philadelphia, Pennsylvania. 471 p. 
Lubinski, K.S., R.E. Sparks, and L.A. Jahn. 1974. The development of 

toxicity indices for assessing the quality of the Illinois River. 

Illinois Water Resources Center Research Report No. 96, University 

of Illinois, Urbana-Champaign, Illinois. 46 p. 
Macaitis, W., J. Variakojis, and R. Kuhl. 1987. Water quality proposal. 

Metropolitan Sanitary District of Greater Chicago. 86 p., 7 

appendices. 
Marvin, D.E., and A.G. Heath. 1968. Cardiac and respiratory responses 

to gradual hypoxia in three ecologically distinct species of 

freshwater fish. Comparative Biochemistry and Physiology 27:349- 

355. 
Mills, H.B., W.C. Starrett, and F.C. Bellrose. 1966. Man's effect on 

the fish and wildlife of the Illinois River. Illinois Natural 

History Survey Biological Notes No. 57. 24 p. 
Merkens, J. C. and K. M. Downing. 1957. The effect of tension of 

dissolved oxygen on the toxicity of un-ionized ammonia to several 

species of fish. Annals of Applied Biology 45:521-527. 
Muchmore, C.B., W.M. Lewis, R.C. Heidinger, M.H. Paller, and L.J. 

Wawronowicz. 1979. Impact of the existing ammonia nitrogen waste 

quality standard. Illinois Institute of Natural Resources Project 

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Toxicity index July 14, 1988 Page 29 

Nos. 80.137, 80.138, and 80.153. 
Neuhold, J. M. and W. F. Sigler. 1960. Effects of sodium fluoride on 

carp and rainbow trout. Transactions of the American Fisheries 

Society 89:358-370. 
Reinbold, K.A., and S.M. Pescitelli. 1981. Effects of cold temperature 

on toxicity of ammonia to rainbow trout, bluegills and fathead 

minnows. Contract Report. Contract No. 68-01-5832. U. S. 

Environmental Protection Agency, Region V, Water Division, Chicago, 

Illinois. 25 p. 
Reinbold, K.A., and S.M. Pescitelli. 1982. Effects of exposure to 

ammonia on sensitive life stages of aquatic organisms. Contract 

Report. Contract No. 68-01-5832. U. S. Environmental Protection 

Agency, Region V, Water Division, Chicago, Illinois. 33 p. 
Roseboom, D.P. and D.L. Richey. 1977. Acute toxicity of residual 

chlorine and ammonia to some native Illinois fishes. Report of 

Investigation 85. Illinois State Water Survey, Urbana, Illinois. 

42 p. 
Sparks, R.E. 1984. The role of contaminants in the decline of the 

Illinois River: implications for the Upper Mississippi. Pages 25-65 

in J.G. Wiener, R.V. Anderson, D.R. McConville, eds. Contaminants in 

the Upper Mississippi River. Butterworth Publishers, Stoneham, 

Massachusetts. 384 p. 
Sparks, R.E. and K.B. Anderson. 1977. Toxicity of ammonia in mixtures 

and development of a toxicity index for use in a stream 

classification system for Illinois: summary report. Pages 34-40 in. 

B.B. Ewing, ed. Feasibility of a systematic approach to water 

quality management in Illinois. Report of the Stream/Lake 

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Toxicity index July 14, 1988 Page 30 

Classification Project. Illinois Institute for Environmental 

Quality Document No. 77/35. 
Sparks, R.E., and W.C. Starrett. 1975. An electrofishing survey of the 

Illinois River, 1959-1974. Illinois Natural History Survey Bulletin 

31(8):317-380. 
Sprague, J. B. and B. A. Ramsay. 1965. Lethal levels of mixed copper- 
zinc solutions for juvenile salmon. Journal Fisheries Research 

Board of Canada 22:425-432. 
Sprague, J.B. 1970. Measurement of pollutant toxicity to fish. 1. 

Bioassay methods for acute toxicity. Water Research 4:3-32. 



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FIGURE LEGENDS AND LIST OF TABLES FOR TOXICITY INDEX MANUSCRIPT BY S 

Figure 1. Mean and maximum toxicity indices in the Illinois Waterway, 1972- 
1974. River mileages begin downstream at the confluence with the 
Mississippi, at river mile 0, and progress upstream toward Chicago 
and Lake Michigan, at river mile 330. The horizontal line is drawn 
at a toxicity index value of 0.2, below which populations of native 
species can maintain themselves, if factors other than acute 
toxicity are not limiting. An index of 1.0 is lethal, equivalent 
to the 96-hour LC50. See text for explanation. 

Figure 2. Cyanide and ammonia component toxicities in the Illinois Waterway, 
1972-1974. River mileages begin downstream at the confluence with 
the Mississippi, at river mile 0, and progress upstream toward 
Chicago and Lake Michigan, at river mile 330. The horizontal line 
is drawn at a toxicity index value of 0.2, below which populations 
of native species can maintain themselves, if factors other than 
acute toxicity are not limiting. An index of 1.0 is lethal, 
equivalent to the 96-hour LC50. See text for explanation. 

Figure 3. Toxicity indices at 1-hour intervals in 3 reaches of the DuPage 
River on 17 March 1972. Indices are output from a simulation 
model, calibrated for the DuPage River. 

Table 1. Toxicity of a total ammonia-N concentration of 47 mg/1 in the 
Illinois River under different conditions. 

Table 2. The number of excursions past different toxicity levels in the 
DuPage River during a simulated 3-year period. 

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Table 3. Impact of a water quality management plan on toxicity in the DuPage 
River. 



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TABLE 1 —Toxicity of a total ammonia-N concentration of 47 mg/1 to 1-gram 
bluegills under different conditions of temperature (°C), pH, and 
dissolved oxygen (DO). 





PH 




DO 


Bluegill Toxicity 


Temp 


mg/1 


% Saturation 


Index 


28 


8.0 


7.5 


100 


2.69 


28 


8.0 


5.0 


66 


4.83 


28 


8.0 


3.0 


40 


13.4 


28 


8.0 


1.0 


13 


3080 


28 


7.0 


5.0 


66 


0.514 


28 


7.5 


5.0 


66 


1.60 


28 


8.0 


5.0 


66 


4.83 


28 


8.5 


5.0 


66 


13.4 


4 


8.0 


12.2 


100 


1.90 


8 


8.0 


11.1 


100 


2.05 


12 


8.0 


10.1 


100 


2.22 


4 


8.0 


8.1 


66 


3.43 


4 


8.0 


4.9 


40 


9.33 


4 


8.0 


1.6 


13 


534 



TABLE 2 — The number of excursions past different toxicity levels 
for various lengths of time during a 3-year period, 
1 October 1970 to 30 September 1973, in 3 reaches of the 
DuPage River (Lubinski 1981). 







Num 


iber 


of To 


xicity 


Index 






Excursions 


for 3 


Duration Times 




Toxicity Index 


























BGTUs 


1 h 




24 h 




96 h 


Fishless reach 


1.0 


73 




51 




42 




2.0 


155 




55 




34 




3.0 


194 




39 




21 


Carp reach 


1.0 


74 




31 




14 




2.0 


16 




1 









3.0 


12 




1 







Bluegill reach 


0.1 


194 




29 




15 




0.2 


37 




25 




11 




0.3 
















TABLE 3 — Impact of a hypothetical water quality management 
plan (see text for explanation) upon the toxicity 
index in the fishless reach, DuPage River, DuPage 
County, Illinois (Brigham and Hey 1981). 





Without 


With 




Plan 


Plan 


Toxicity Index 






max. 


784 


51.5 


min. 


0.120 


0.118 


mean. 


23.0 


2.12 


ComDonent Toxicity 






Ammonia 


20.2 


0.023 


Cyanide 


0.088 


0.079 


Lead 


0.023 


0.000 


Zinc 


0.004 


0.003 


Copper 


0.020 


0.008 


LAS 


0.772 


0.073 


Chlorine 


J. 94 


1.94 



Illinois Waterway 
Mean Toxicity Indices 




River Mile 



Illinois Waterway 
Maximum Toxicity Indices 



1972 




340 



River Mile 



Illinois Waterway 
Cyanide Component Toxicity 




340 



0.30 - 



C 0.20 



Illinois Waterway 
Ammonia Component Toxicity 




River Mile 



CO 




13 


CM 


E 


N- 


'(f) 


o> 


5— 


SI 





o 


> 


CO 


be 


^ 





h- 


D) 




CC5 




a. 




13 




Q 






CO 

Q 





O 
o 

X 



s^iun Ajiojxoi ||iBen|g 



Literature Cited 



Cairns, J., Jr., R.E. Sparks, and W.T. Waller. 1973. A tentative proposal 
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528. 256 pp. 

Dickson, K.L., D. Gruber, C. King, and K. Lubenski. 1980. Biological 

monitoring to provide an early warning of environmental contaminants. 
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Hughes, R.M., D.P. Larsen, and J.M. Omernik. 1986. Regional reference sites: 
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Hughes, R.M., E. Rexstad, and C.E. Bond. 1987. The relationship of aquatic 
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Larsen, D.P., R.M. Hughes, J.M. Omernik, D.R. Dudley, CM. Rohm, R.T. 
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Mills, H.B., W.C. Starrett, and F.C. Bellrose. 1966. Man's effect on the 
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